Pharmaceuticals, Personal Care Products, and Endocrine...

ENVIRONMENTAL ENGINEERING SCIENCEVolume 20, Number 5, 2003© Mary Ann Liebert, Inc.

Pharmaceuticals, Personal Care Products, and EndocrineDisruptors in Water: Implications for the Water Industry

Shane A. Snyder,1,* Paul Westerhoff,2 Yeomin Yoon,2 and David L. Sedlak3

1Department of Research and Development Southern Nevada Water Authority

Boulder City, NV 890052Department of Civil and Environmental Engineering

Arizona State UniversityTempe, AZ 85287

3Department of Civil and Environmental EngineeringUniversity of California

Berkeley, CA 94720

ABSTRACT

For over 70 years, scientists have reported that certain synthetic and natural compounds could mimic nat-ural hormones in the endocrine systems of animals. These substances are now collectively known as en-docrine-disrupting compounds (EDCs), and have been linked to a variety of adverse effects in both humansand wildlife. More recently, pharmaceuticals and personal care products (PPCPs) have been discovered invarious surface and ground waters, some of which have been linked to ecological impacts at trace con-centrations. The majority of EDCs and PPCPs are more polar than traditional contaminants and severalhave acidic or basic functional groups. These properties, coupled with occurrence at trace levels (i.e., ,1mg/L), create unique challenges for both removal processes and analytical detection. Reports of EDCs andPPCPs in water have raised substantial concern among the public and regulatory agencies; however, verylittle is known about the fate of these compounds during drinking and wastewater treatment. Numerousstudies have shown that conventional drinking and wastewater treatment plants can not completely removemany EDCs and PPCPs. Oxidation with chlorine and ozone can result in transformation of some com-pounds with reactive functional groups under the conditions employed in water and wastewater treatmentplants. Advanced treatment technologies, such as activated carbon and reverse osmosis, appear viable forthe removal of many trace contaminants including EDCs and PPCPs. Future research needs include moredetailed fate and transport data, standardized analytical methodology, predictive models, removal kinetics,and determination of the toxicological relevance of trace levels of EDCs and PPCPs in water.

Key words: endocrine disruptor; pharmaceutical; drinking water; wastewater; treatment; review

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ENDOCRINE DISRUPTING COMPOUNDS

PERHAPS THE MOST DIFFICULT PART of understanding thesubject of endocrine disruption involves a definition

of the term. The Environmental Protection Agency (EPA)has defined environmental endocrine disrupting com-pounds (EDCs) as exogenous agents that interfere with the“synthesis, secretion, transport, binding, action, or elimi-

*Corresponding author: Department of Research and Development, Southern Nevada Water Authority, 243 Lakeshore Road,Boulder City, NV 89005. Phone: 702-567-2317; Fax: 702-567-2085; E-mail: [email protected]

nation of natural hormones in the body that are responsi-ble for the maintenance of homeostasis, reproduction, de-velopment, and/or behavior” (EPA, 1997). However, def-initions and opinions on what defines an EDC vary greatly.It is generally accepted that the three major classes of en-docrine disruption endpoints are estrogenic (compoundsthat mimic or block natural estrogen), androgenic (com-pounds that mimic or block natural testosterone), and thy-roidal (compounds with direct or indirect impacts to thethyroid). As we will illustrate, the majority of research thusfar has focused only on estrogenic compounds; however,disruption of androgen and thyroid function may be ofgreater or equal importance biologically.

Although the topic of endocrine disruption is consid-ered an “emerging issue” in the water industry, scientistshave known about the ability of natural and synthetic com-pounds to interfere with the hormone systems of animalsfor over 70 years. The discovery that certain compoundscan mimic the endogenous hormones of animals was re-ported as early as the 1930s (Cook et al., 1934; Walkerand Janney, 1930). In 1940, Stroud reported that certainsynthetic chemicals were estrogenic (Stroud, 1940). Anarticle from the journal Science published in 1946 ex-plained that the molecular configurations of natural and

synthetic compounds influenced the degree of estrogenicand androgenic bioactivity in rodents (Schueler, 1946).The ability of estrogenic and androgenic compounds tointerfere with the natural metamorphosis of amphibianswas reported as early as 1948 (Sluczewski and Roth,1948). These early papers began to describe how the mo-lecular configurations of natural and synthetic compoundscould mimic the primary endogenous female hormone17b-estradiol (E2) and the male hormone testosterone.More importantly, these early studies laid the foundationfor how these hormone-mimicking compounds could re-sult in reproductive toxicity. Figure 1 provides the struc-tures for some estrogenic compounds.

The estrogenic activity of synthetic organic compoundswas of little interest to the environmental community untilthe discovery that the organochlorine pesticide, DDT andits metabolites had endocrine-disrupting properties (Fisheret al., 1952; Bitman et al., 1968; Welch et al., 1969; Bit-man and Cecil, 1970; Wrenn et al., 1970). The connectionbetween endocrine disruption and reproductive failures inwildlife was not made until the 1980s, when it was reportedthat gulls living in areas contaminated with DDT exhibiteddeformed sex organs and skewed sex ratios (Fry and Toone,1981; Fry et al., 1987). This was one of the first docu-

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Figure 1. Structures of estrogenic compounds.

mented connections between an environmental contami-nant and reproductive impacts via a hormone-mediatedmechanism. Possible links between organochlorine pesti-cides, including DDT and its metabolites, and endocrinedisruption were provided by researchers in Florida, whodiscovered reproductive disorders in alligators in LakeApopka (Guillette et al., 1994, 1996).

More recent studies have further demonstrated the im-portance of endocrine disruption in wildlife populations.For example, marine gastropods exposed to tributyltins,which leach from certain antifouling paints and PVCpipes, experienced severe population declines and repro-ductive disorders including imposex (development ofmale sex characteristics in females) (Gibbs et al., 1991).In some amphibian populations, supernumerary limbsand missing limbs have been attributed to certain pesti-cides and other anthropogenic chemicals (Ouellet et al.,1997; Sparling, 2000). In particular, trace concentrationsof the widely used herbicide atrazine have been associ-ated with endocrine disruption in frogs from the Mid-western United States (Hayes et al., 2002). Degradationproducts from the widely used alkylphenol polyethoxy-late (APE) surfactants, which are ubiquitous contami-nants of wastewater treatment plant effluents, have beenshown to be estrogenic (Mueller and Kim 1978; Whiteet al, 1994; Routledge and Sumpter 1997; Giesy et al.,2000) and bioaccumulative (Ahel et al., 1993; Liber etal., 1999; Lye et al., 1999; Snyder et al., 2001a). In the1990s, reports from the United Kingdom and the UnitedStates indicated that fish living below wastewater treat-ment plants had several reproductive abnormalities (Be-vans et al., 1996; Folmar et al., 1996; Harries et al., 1996;Purdom et al., 1994; Jobling et al., 1998). These repro-ductive abnormalities included changes in the levels ofsex steroids, gonadal histology (e.g., hemaphrodism andintersex), and increased levels of the female egg yolk pre-cursor, vitellogenin, in male fish. Collectively, these im-pacts of wastewater effluent on male fish are referred toas feminization because fish that are genetically male ex-hibit female sex characteristics. However, conclusivedata linking population level impacts on fish from EDCexposures has not yet been published. There are manyother examples of exogenous compounds acting as EDCs,and more EDCs will likely be discovered as screeningand testing methods become available.

Endocrine disruption also can be caused by naturallyoccurring compounds. For example, estrogens from plantsources, known as phytoestrogens, have been linked toreproductive failures in animals since the 1930s (Walkerand Janney, 1930; Levin et al., 1951; Brookbanks et al.,1969; Metzler and Pfeiffer, 1995; Safe and Gaido, 1998).This effect was evident in sheep grazing on certain strainsof clover in New Zealand. These sheep exhibited severe

reproductive impairment due to phytoestrogens(Millington et al., 1964; Adams, 1998). Likewise, theinability of captive cheetahs to reproduce at the Cincin-nati Zoo was linked to a diet high in phytoestrogens(Setchell et al., 1987). In 1951, a vegetable oil was foundto contain various phytoestrogens and phytoandrogens(Levin et al., 1951). Interestingly, various over-the-counter medicinal supplements, such as those recom-mended for estrogen replacement therapy in postmeno-pausal women, contain high levels of phytoestrogens.Research is necessary to assess the relative importanceof phytoestrogens in human diet, both by direct inges-tion of vegetables and vegetable products and possiblyfrom bioaccumulation of phytoestrogens in meat prod-ucts. Likewise, industrial activities that release largequantities of phytoestrogens may have adverse effectson aquatic ecosystems as evidenced by recent studiesdocumenting masculinization of fish exposed to effluentfrom pulp and paper mills and the presence of andro-genic compounds in these effluents (Munkittrick et al.,1997; Bortone and Cody, 1999; Larsson et al., 2000;Jenkins et al., 2001). Likewise, the degradation of veg-etable matter and paper products in wastewater treatmentplants (WWTPs) may contribute to releases of phytoe-strogens into the aquatic ecosystem.

Initial attempts to identify the cause of feminization offish exposed to wastewater effluent focused on syntheticorganic chemicals with known estrogenic effects, such asplasticizers and APE surfactant degradation products.However, a series of studies employing in vitro bioassay-directed chemical fractionation implicated the human hor-mone, 17b-estradiol (E2), and the synthetic birth controlpharmaceutical, 17a-ethinyl estradiol (EE2), as the mostpotent estrogens in these complex mixtures (Desbrow etal., 1998; Snyder et al., 1999, 2001c). Related researchinvolving exposure of fish to E2 and EE2 under labora-tory conditions at concentrations as low as 2 ng/L couldinduce measurable changes in fish reproduction (Arcand-Hoy et al., 1998; Kramer et al., 1998; Panter et al., 1998;Routledge et al., 1998). These reports indicated that apharmaceutical (EE2) could induce endocrine-disruptiveeffects in fish at concentrations present in some munici-pal WWTP effluents. This cause–effect relationship hasstimulated a great deal of new research on the identifica-tion of trace pharmaceuticals in the environment.

The unexpected impacts of trace concentrations ofEDCs on wildlife raised concerns about the potential ef-fects of these chemicals on humans (Colborn et al., 1997).The best documented instance of endocrine disruption inhumans involved in utero exposure to the potent syntheticestrogen diethylstilbestrol (DES), which resulted in ad-verse reproductive impacts in human offspring (Herbst etal., 1971; Gill et al., 1979). In this case, DES exposure

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was not as an environmental contaminant; it was used as a pharmaceutical administered to pregnant women(Sower et al., 1983; Rumsey and Hammond, 1990). How-ever, the unanticipated sensitivity of the developing hu-man reproductive system to DES clearly demonstratesthat the human embryo/fetus is not immune to insult byexogenous chemicals that act as hormones. Besides us-age as a human pharmaceutical, DES is also used for cer-tain agricultural applications such as increasing livestockgrowth and milk production.

Some researchers have attributed decreases in humansperm quality and quantity over the past 5 decades to endocrine disrupting compounds in the environment(Sharpe and Skakkebaek, 1993; Stone, 1994; Carlsen etal., 1995). Likewise, it has been suggested that sharp in-creases in breast, testicular, and prostate cancers reportedover the past 40 years are related to endocrine disruptingcompounds in the environment (Krishnan and Safe, 1993;Ahlborg et al., 1995; Carlsen et al., 1995; Ashby, 1997;EPA, 1997; Gillesby and Zacharewski, 1998). However,this topic is still quite controversial, and other scientistshave produced data refuting these arguments. If EDCsare causing adverse human health effects, it is unlikelythat these effects would be caused by estrogenic chemi-cals in water due to their extremely minute concentra-tions that result in doses that are small compared to phy-toestrogens and other estrogenic compounds present infood sources. Estrogenic hormones in water are less likelyto cause adverse effects in humans than they are in fishdue to differences in exposures. Although fish may beconstantly exposed to EDCs present in the aquatic envi-ronment, humans are exposed mainly through ingestionof limited quantities of water.

In addition to compounds capable of eliciting estro-genic activity, concerns have been raised regarding hu-man health effects associated with pollutants that inter-act with other hormone systems. Recently, a great dealof public and regulatory interest has focused on perchlo-rate (ClO4

2), which affects the thyroid gland by com-petitively inhibiting iodide transport (Urbansky, 2000;Logan, 2001). Ammonium perchlorate is a strong oxi-dizer that has been used extensively in solid-rocket fu-els, fireworks, matches, and other industrial and domes-tic uses (Urbansky, 2000). Nearly every state in theUnited States has known perchlorate use, storage, ormanufacturing sites, and perchlorate contamination hasrecently been found in many ground and surface waters(Wang et al., 2002). Perchlorate salts were also used asa pharmaceutical to treat overactive thyroid disease andcontinue to be used to treat side effects of certain che-motherapy drugs (Wang et al., 2002). Because perchlo-rate has a direct impact on the thyroid, it is consideredan EDC, and will likely become the first drinking water

contaminant regulated on endocrine disrupting toxicity,as the U.S. EPA is currently establishing a reference dose.

Some scientists have suggested that some drinking wa-ter disinfection byproducts (DBPs) may act as EDCs.Several reports, which include epidemiological studies,have reported increases in spontaneous abortions andcancers in humans as linked to elevated concentrationsof halogenated disinfection byproducts (King and Mar-rett, 1996; King et al., 2000a, 2000b). Additionally, theEndocrine Disruptor Screening and Testing AdvisoryCommittee (EDSTAC), a Federal Advisory Committee,suggested that DBPs should be included among the mix-tures to be evaluated for endocrine disruptive effects(EPA, 1998). Because DBPs generally are orders of mag-nitude greater in concentration than EDCs and pharma-ceutical and personal care products (PPCPs), and havebeen implicated as human reproductive toxicants, effortsto control EDCs and PPCPs by oxidation ultimately mayprove to be counterproductive due to increased byprod-uct formation potential.

PHARMACEUTICALS AND PERSONALCARE PRODUCTS (PPCPs)

The first concerns regarding potential adverse effectsof pharmaceuticals in municipal wastewater were ex-pressed by Stumm-Zollinger and Fair in 1965 and Tabakand Bunch in 1970 (Stumm-Zollinger and Fair, 1965;Tabak and Bunch, 1970). In their visionary, but largelyforgotten study of the biotransformation of estrogenichormones by activated sludge, they noted that natural andsynthetic estrogens could pose an ecological threat. Otherstudies conducted in the United States during the 1970sdocumented the presence of other pharmaceuticals, suchas clofibric acid and salicylic acid in wastewater efflu-ents (Garrison et al., 1975; Hignite and Azarnoff, 1977).Later studies have shown that natural and synthetic es-trogens are ubiquitous contaminants of wastewater attrace concentrations (Tabak et al., 1981; Aherne et al.,1985; Desbrow et al., 1998; Lee and Peart, 1998; Sny-der et al., 1999; Huang and Sedlak, 2001). However, re-ports demonstrating the possible impacts of trace levels(ng/L) of the pharmaceutical EE2 (Arcand-Hoy et al.,1998; Desbrow et al., 1989; Renner, 1998; Snyder et al.,2001c) have stimulated a dramatic increase in researchon these compounds as environmental contaminants(Halling-Sorensen et al., 1998; Daughton and Ternes,1999; Snyder et al., 2001a). A wide variety of PPCPshave now been reported as environmental contaminantsincluding antibiotics, X-ray contrast media, analgesics,antiseptics, and many others (Halling-Sorensen et al.,1998; Daughton and Ternes, 1999). In the United States,

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a recent survey indicated widespread PPCP contamina-tion of streams (Kolpin et al., 2002). Table 1 shows sev-eral of the most frequently detected EDCs and PPCPsalong with concentration, usage, and structural informa-tion. Unfortunately, sparse data exist to explain the tox-icological relevance of trace pharmaceutical compounds.Nevertheless, public perception regarding the presence ofPPCPs in water supplies has increased attention to thisissue despite the very low concentrations reported.

U.S. REGULATORY ISSUES

In the United States, regulation of contaminants indrinking water began in 1962, with the Public HealthServices Standards. These regulations included severalcompounds now known to be endocrine disruptors, suchas arsenic, cadmium, and some phenols. The SafeDrinking Water Act of 1974 is the principal law gov-erning drinking water safety in the United States. Thislaw required the EPA to establish maximum contami-

nant levels for various drinking water contaminants in-cluding some pesticides now known to have endocrine-disruptive activity. However, endocrine disruption wasnot specifically named in any United States legislationuntil 1995 with amendments to the Safe Drinking Wa-ter Act (bill number S.1316) and Food Quality Protec-tion Act (bill number P.L. 104-170), mandating thatchemicals and formulations be screened for potential en-docrine activity before they are manufactured or usedin certain processes where drinking water and/or foodcould become contaminated. Under these laws, the EPAis required to “develop a screening program, using ap-propriate validated test systems and other scientificallyrelevant information, to determine whether certain sub-stances may have an effect in humans that is similar toan effect produced by a naturally occurring estrogen, orother such endocrine effect as the Administrator maydesignate.” Furthermore, these laws specify that theEPA must have developed a testing program by 1998,implemented the program by 1999, and reported to Con-gress by 2000.



Table 1. Compounds with highest frequency of detection in recent USGS EDC/PPCP survey of U.S. streams (Kolpin et al., 2002).

Frequency of detection LogKow,measuredCompounds Use (%) (calculated)

Coprostanol Estrogen ,80% (8.82)Cholesterol Plant/animal steroid ,80% (8.74)N-N-diethyltoluamide Mosquito repellant ,80% 2.18 (2.26)Caffeine Stimulant ,75% 20.07 (0.16)Tris(2-chloroethyl)phosphate Fire retardant ,75% 1.44 (1.63)Triclosan Antibiotic ,60% NA4-Nonylphenol Surfactant ,60% (5.92)4-Nonylphenol monoethoxylate Surfactant ,50% NAEthanol, 2-butoxy-phosphate Plasticizer ,45% NA4-Octylphenol monoethoxylate Surfactant ,45% NABisphenol A Plasticizer ,45% 3.32 (3.64)Cotinine Nicotine metabolite ,35% 0.07 (0.34)4-Nonylphenol diethoxylate Surfactant ,35% NA5-Methyl-1H-benzotriazole Antioxidant ,30% NAFluoranthene PAH ,30% 5.16 (4.93)1,7,-Dimethylxanthine Caffeine metabolite ,30% 20.22 (20.39)Pyrene PAH ,25% 4.88 (4.93)Trimethoprim Antibiotic ,25% NA1,4-Dichlorobenzene Deodorizer ,25% 3.44 (3.28)Acetaminophen Analgesic ,25% 0.46 (0.27)Tetrachloroethylene Solvent ,20% NA4-Octylphenol diethoxylate Surfactant ,20% NAErythromycin-H2O Antibiotic ,20% NAEstriol Estrogen ,20% 2.45 (2.81)Lincomycin Antibiotic ,15% 0.59 (0.29)Sulfamethoxazole Antibiotic ,15% 0.89 (0.48)Phthalic anhydride Plasticizer ,15% 1.60 (2.07)Carbaryl Insecticide ,15% NA

To meet the requirements of this recent legislation, theEPA formed the EDSTAC to provide recommendationson a conceptual framework, priority setting, screening,and testing methodologies, and communication and out-reach programs. The EDSTAC group consisted of vari-ous stakeholders and experts in reproductive toxicology.The committee began deliberations in October of 1996and issued a final report in July of 1998 recommendingthat human and wildlife impacts be considered, and thatestrogen, androgen, and thyroid (EAT) end points be ex-amined (EPA, 1998). The conceptual framework devisedby EDSTAC consists of an initial sorting, prioritization,Tier 1 and 2 testing, and a hazard assessment of an esti-mated 87,000 chemicals. In addition to discrete chemi-cals, EDSTAC recommended the evaluation of mixturesof chemicals in breast milk, baby formulas, hazardouswaste sites, pesticides and fertilizers, drinking water DBPs,and gasoline.

In 2001, the EPA formed the Endocrine DisruptorMethods Validation Subcommittee (EDMVS) to evalu-ate the test battery suggested by EDSTAC. The EDMVSis tasked with method validation by determining if a par-ticular method is transferable to other laboratories, canbe validated with representative chemicals, has sufficientsensitivity to EAT end points, and has appropriate stan-dard operating procedures. Some difficult issues en-countered in the standardization of EDC testing includeanimal diets, dosing methods and ranges, testing of mix-tures, and interspecies comparisons. The outcome of thisscreening battery is critical to the water industry, as it de-signed to definitively identify EDCs. However, it is im-portant to note that the current legislation regulates onlythe industries producing or using raw chemicals, and notthe water industry. As a result, these actions may havelittle immediate effect on water and wastewater treatmentregulations.

There are currently no federal regulations for pharma-ceuticals in drinking or natural waters. The Food andDrug Administration (FDA) requires ecological testingand evaluation of a pharmaceutical only if an environ-mental concentration in water or soil is expected to ex-ceed 1 mg/L or 100 mg/kg, respectively (FDA, 1998). Inlight of the recent data on the occurrence of PPCPs in theaquatic environment, these policies may need to be re-considered. Although extensive monitoring programs areunderway, toxicological studies conducted at environ-mentally relevant concentrations are necessary for intel-ligible regulations to be established. The State of Cali-fornia has been considering the potential impacts ofEDCs and PPCPs, especially when municipal wastewatereffluent is used for indirect potable reuse. A recent mod-ification to California’s draft regulations for indirectpotable reuse states, “Each year, the PGRRP [planned

groundwater recharge reuse project] shall monitor the re-cycled water for endocrine disrupting chemicals andpharmaceuticals specified by the Department, based on areview of the PGRRP engineering report and the affectedgroundwater basin(s).” Although the regulations have notbeen finalized, many practitioners of indirect potablereuse in California are establishing monitoring programfor EDCs and PPCPs. Because California’s water reuseprogram often establishes precedents for programsthroughout the world, it is likely that other regulatoryagencies will adopt similar language in their own waterrecycling programs.

ANALYTICAL METHODS

Because EDCs represent a broad variety of com-pounds, it is important to define which EDCs one seeksto analyze. It is widely accepted that DDT and otherorganochlorine pesticides can act as EDCs. Methodolo-gies for these “classic” contaminants, as well as variousendocrine-disrupting metals, are well established withstandardized protocols used in drinking and wastewaterregulations. The majority of novel analytical work is fo-cused on trace levels of less-characterized contaminantswith greater polarity than many of the “classic” contam-inants. Several classes of EDCs and PPCPs have acidicor basic moieties, large molecular weights, and/or polarfunctional groups. No standard methods are currentlyavailable for these compounds, and few commercial laboratories analyze these compounds. A further compli-cation is the desire to quantitate these compounds at ultratrace concentrations (sub-ng/L), which may havetoxicological relevance (e.g., EE2).

A vast array of analytical methodologies has been ap-plied for the quantitation of EDCs and PPCPs in water.Although direct measurement is possible (Yoon et al.,2003), the majority of methods involve an extraction pro-cedure followed by instrumental and/or immunoassayanalyses (Tabak and Bunch, 1970; Keith et al., 1975;Hignite and Azarnoff, 1977; Aherne et al., 1985; Shoreet al., 1993; Ternes et al., 1998; Snyder et al., 1999,2001a; Stumpf et al., 1999; Heberer and Dumnbier, 2000;Huang and Sedlak, 2001). The type of extraction, vol-ume of water used, and type of analytical equipment useddepend on the compounds to be analyzed and source ofwater. For instance, WWTP effluents are the majorsource of these compounds to surface waters. The anal-ysis of WWTP effluents provides unique challenges fortrace quantification because the effluent also contains nu-merous interfering compounds associated with organicmatter. Consequently, extensive extraction, cleanup, andsophisticated instrumentation usually is required to ana-

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lyze these compounds. Several methodologies for quan-titation of EDCs and PPCPs in natural waters used solid-phase extractions followed by instrumental analyses us-ing gas chromatography coupled with mass spectrometricdetection, liquid chromatography with mass spectromet-ric detection, immunoassays, or a combination of tech-niques (Keith et al., 1975; Hignite and Azarnoff, 1977;Tabak et al., 1981; Aherne et al., 1985; Shore et al., 1993;Ternes et al., 1998; Snyder et al., 1999, 2001a; Stumpfet al., 1999; Heberer and Dumnbier, 2000; Huang andSedlak, 2001; Ternes, 2001). Each method has advan-tages and disadvantages, detailed descriptions of whichare outside the scope of this paper.

Bioanalytical techniques are important tools for mon-itoring certain EDCs and PPCPs. These techniques em-ploy a biological end point that is related to a type oftoxicity or a class of compounds. The most simple meth-ods are receptor binding assays and cellular bioassaysthat have rapid response times, high sensitivity, and rel-ativity low cost (Welch et al., 1969; Mueller and Kim,1978; Anderson et al., 1996; Routledge and Sumpter,1997; Zacharewski, 1997; Snyder et al., 2000, 2001c).In vivo bioassays are also used to detect various classesof EDCs. The most common biomarker for estrogenicexposure in the aquatic environment is an increase inplasma vitellogenin (an egg yolk lipid) in male fish. Fishcan be caged in various susceptible waters and tested forvarious endocrine biomarkers to determine the extent ofEDC exposure. Several studies of this nature have foundsignificant reproductive impacts in fish caged belowwastewater outfalls (Purdom et al., 1994; Rudel, 1997;Routledge et al., 1998; Miles-Richardson et al., 1999;Snyder et al., 2000). Although each instrumental andbioanalytical technique has advantages and disadvan-tages, a combination of these methods is most likely todetect and quantify EDCs. This approach is often re-ferred to as toxicity identification evaluation, and mayuse bioassay-directed fractionation to guide instrumen-tal analyses towards elucidation of toxic compounds(Desbrow et al., 1998; Routledge et al., 1998; Castilloand Barcelo, 1999; Snyder et al., 2001c). The issue ofmixture toxicity, especially in relation to endocrine dis-ruption, may likely drive the drinking water industry toemploy a number of biological monitoring tools. Sinceall environmental contaminants exist as components ofcomplex mixtures, biological tests may be the only wayto assess the total endocrine-disruptive potential of con-taminants in water.

EDC/PPCP removal during water treatment

EDCs and PPCPs in municipal wastewater can be re-moved during wastewater treatment, during their pas-sage through surface or groundwater (e.g., natural at-

tenuation) or during drinking water treatment. To assessthe potential for exposure of humans and aquatic or-ganisms to these compounds and to design more effec-tive treatment systems, environmental professionalsneed to understand the mechanisms through whichEDCs and PPCPs are attenuated in engineered and nat-ural systems. The best understood of the treatment pro-cesses is conventional wastewater treatment (Ternes,1998; Ternes et al., 1999a, 1999b; Johnson andSumpter, 2001). Compared to municipal wastewatertreatment plants, much less is known about the behav-ior of EDCs and PPCPs in receiving waters and in drink-ing water treatment plants. Therefore, the purpose ofthis review is to summarize existing data, to predict thebehavior of different classes of EDCs and PPCPs basedupon chemical properties, and to identify areas whereadditional research is needed.

Conventional surface water treatment plants (SWTPs)typically treat water via coagulation using alum, ferricchloride, and/or synthetic polymers followed by floccu-lation, sedimentation, filtration, and disinfection. Con-ventional SWTPs achieve high removals of pathogensand other biological particles, and modest removal of dis-solved organic carbon (DOC, 1 to 10 mg/L). In the UnitedStates, chlorine and chloramines are typically used fordisinfection, while ozonation is more commonly prac-ticed in European countries. Alternative SWTP designsor modifications to conventional WTPs sometimes in-clude additional water treatment processes (e.g., activatedcarbon, biofiltration, membranes, aeration, chemical soft-ening, ultraviolet light irradiation).

Our understanding of the removal of EDCs and PPCPsin receiving waters and in drinking water treatment sys-tems is limited because the analyses for these compoundsare rare, and when detected, they are present at fluctuat-ing concentrations near analytical method detection lim-its. With a few notable exceptions, most of our knowl-edge about the removal of these compounds is derivedfrom laboratory or bench-scale studies. When data on re-moval of EDCs and PPCPs are not available, it may bepossible to make predictions based upon results of previous research with contaminants exhibiting similarchemical properties.

As a result of much evidence linking estrogenic hor-mones to endocrine disruption in fish, there has been agreat deal of interest in the fate of E2, EE2, and nonylphe-nol ethoxylates (NPEs) in surface waters. For example,laboratory studies designed to simulate microbial trans-formation of NPEs in surface waters indicate that theyare degraded in 4 days (Mann and Boddy, 2000), whichis consistent with the observation of a relatively small de-crease in concentrations when nitrified wastewater efflu-ent passes through an engineered treatment wetland.



Studies conducted at a drinking water treatment plantalong the Severn Trent Region, a system with significantinputs of wastewater effluent, indicate that estrogenichormones and other compounds capable of producing es-trogenic effects in bioassays are removed during drink-ing water treatment (Fawell et al., 2001).

The attenuation of PPCPs in receiving waters and indrinking water treatment plants also has been demon-strated in several studies and has been reviewed recently(Heberer, 2002). For example, the attenuation of severalPPCPs was observed as secondary wastewater effluentwas infiltrated into groundwater (Drewes et al., 2001a)and during river bank filtration (Heberer, 2002). How-ever, certain compounds, such as the iodinated X-ray con-trast media were not removed during groundwater infil-tration (Putschew et al., 2000). In contrast to the efficientremoval of many PPCPs observed in saturated and un-saturated groundwater, conventional drinking water treat-ment plants do not appear to remove PPCPs well. For ex-ample, clofibric acid has been detected at concentrationsas high as 270 ng/L in drinking water from Berlin(Heberer and Dumnbier, 2000). Of 47 wastewater trac-ers and EDCs analyzed, 15 were detected in raw drink-ing water (river water) samples, and 14 in finished drink-ing water samples from Atlanta (Henderson et al., 2001).In that study, caffeine was present in all raw waters andsome finished waters. Although all at trace concentra-

tions (ng/L), the following compounds were detected infinished drinking waters: tri(2-chloroethyl)phosphate,phthalic anhydride, triclosan, fluoranthene, pyrene, 2,6-d-t-butylphenol, thanol-2-butoxy-phosphate, and tributylphosphate. Caffeine, cotinine, and acetaminophen werealso detected in the finished drinking water (Hendersonet al., 2001).

Although monitoring studies involving low concentra-tions of EDCs and PPCPs provide insight into the oc-currence of these compounds in the environment, previ-ously discussed issues associated with method detectionlimits and dilution factors complicate the analysis ofmonitoring data. As a result, more insight into removalprocesses can be obtained by examining specific treat-ment processes. Table 2 represents anticipated perfor-mance of different unit processes based upon literaturereports with specific classes of compounds or compoundswith similarities to other trace pollutants that have beenstudied in more detail. The following sections describethe potential for common water treatment processes toremove EDCs or PPCPs.

Chemical precipitation processes

Metal salts (aluminum sulfate, ferric chloride) andsoftening chemicals (calcium oxide, sodium carbonate)are commonly added to destabilize particles present in

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Table 2. Unit processes and operations used for EDCs and PPCPs removal.

Coagulation/ Softening/ DegradationGroup Classification AC BAC O3 /AOPs UV Cl2 /ClO 2 flocculation metal oxides NF RO {B/P/AS}a

EDCs Pesticides E E L-E E P-E P G G E E {P}Industrial chemicals E E F-G E P P-L P-L E E G-E {B}

Steroids E E E E E P P-L G E L-E {B}Metals G G P P P F-G F-G G E P {B}, E {AS}

Inorganics P-L F P P P P G G E P-LOrganometallics G-E G-E L-E F-G P-F P-L P-L G-E E L-E

PhACs Antibiotics F-G E L-E F-G P-G P-L P-L E E E {B}G-E {P}

Antidepressants G-E G-E L-E F-G P-F P-L P-L G-E E G-EAnti-inflammatory E G-E E E P-F P P-L G-E E E {B}

Lipid regulators E E E F-G P-F P P-L G-E E P {B}X-ray contrast media G-E G-E L-E F-G P-F P-L P-L G-E E E {B and P}

Psychiatric control G-E G-E L-E F-G P-F P-L P-L G-E E G-E

PCPs Synthetic musks G-E G-E L-E E P-F P-L P-L G-E E E {B}Sunscreens G-E G-E L-E F-G P-F P-L P-L G-E E G-E

Antimicrobials G-E G-E L-E F-G P-F P-L P-L G-E E F {P}Surfactants/detergents E E F-G F-G P P-L P-L E E L-E {B}

aB, biodegradation; P, photodegradation (solar); AS, activated sludge; E, excellent (.90%); G, good (70–90%); F, fair(40–70%); L, low (20–40%); P, poor (,20%).

water or to precipitate new particles (coagulation), ag-gregate particles (flocculation), and improve settlingcharacteristics of particles (clarification). Sand filtrationis commonly used after clarification for additional parti-cle removal. Natural organic matter (NOM) and EDCs orPPCPs may adsorb to particles in water and metal hy-droxide particles formed during coagulation. Further-more, chemical precipitation can remove moderately hy-drophobic organic contaminants that have a strongaffinity for adsorbed NOM (Rebhun and Lurie, 1993).However, very little is known about the association ofEDCs or PPCPs to particles present in water treatmentsystems. Therefore, it is difficult to make a priori pre-dictions about the removal of EDCs or PPCPs duringchemical precipitation processes.

Despite these potential shortcomings, considerable in-sight can be obtained by considering the behavior of pes-ticides, herbicides, and polycyclic aromatic hydrocarbons(PAHs) during chemical precipitation processes (Eldiband Aly, 1977; Rebhun et al., 1998). Partitioning of or-ganic compounds onto particles can occur through sev-eral mechanisms. For hydrophobic compounds, parti-tioning can be predicted from octanol–water partitioncoefficients (KOW) (Karickhoff and Morris, 1985; Chiouet al., 1998). Under the conditions encountered duringwater treatment, only those compounds with relativelyhigh KOW values (i.e., .105 will be removed to an ap-preciable degree. As indicated in Table 1, most com-pounds of potential concern are relatively polar (log KOW

values less than 3) and as a result, only a few EDCs andPPCPs (e.g., nonylphenol, fluroanthene, pyrene) are ex-pected to be removed during chemical precipitation.Some of the more hydrophobic compounds are presentin wastewater effluents, but may not occur in raw drink-ing water supplies (Henderson et al., 2001). Some surfactants have relatively high KOW values, but includeboth polar and nonpolar moieties. The partitioning ofoctylphenol in English river sediments was greatest to theclay and silt fraction of the sediments, suggesting a hy-drophobic interaction with organic carbon as well as sur-face area associated with the clay and silt particles (John-son et al., 1998). As a result, only a few compounds (e.g.,bisphenol A, E2, EE2, octylphenol, PAHs) could be as-sociated with organic phases of particles in drinking wa-ter treatment plants.

In addition to hydrophobic partitioning, organic con-taminants can be adsorbed to particles by interactions ofpolar functional groups with charged particles and min-eral surfaces by complexation or ion exchange. For ex-ample, the observed sorption of many polar veterinarypharmaceuticals (e.g., tetracycline) to soil and sedimentis considerably stronger than predicted by hydrophobicinteractions alone (Tolls, 2001). Such interactions could

be particularly important in drinking water treatment,where mineral oxides provide a relatively high density ofsurface functional groups that could interact with polarpharmaceuticals. For example, approximately 25% of theNPEs in groundwater were removed during coagulationwith alum (Fielding et al., 1998). Therefore, it is possi-ble that some polar EDCs or PPCPs could be removedduring coagulation.

Despite the predictions, available data from pilot andfull-scale water treatment plants indicate that removalwill be modest at best. Neither lime softening nor alumcoagulation (conventional or enhanced dosages rang-ing from 6 to 18 mg/L) demonstrated atrazine removal(Zhang and Emary, 1999). Coagulation/flocculation/ sed-imentation with alum and iron salts or excess lime/sodaash softening did not result in significant removal ofantibiotics (i.e., carbadox, sulfachlorpyridazine, sul-fadimethoxine, sulfamerazine, sulfamethazine, sul-fathiazole, and trimethoprim) (Adams et al., 2002). Inanother study, ferric chloride precipitation did not re-move several pharmaceuticals frequently detected insurface waters (diclofenac, carbamazepine, bezafibrate,and clofibric acid) (Sacher et al., 2000). Certain pesti-cides were poorly removed by coagulation and ,50%of the PAHs pyrene, fluoranthene, and anthracene wereremoved through hydrophobic interactions (Eldib andAly, 1977; Rebhun et al., 1998). On the basis of pre-dictions of hydrophobic interactions and results of full-scale measurements, we conclude that EDCs andPPCPs not associated with colloidal or particulate ma-terial will most likely be poorly removed during coag-ulation.

Activated carbon adsorption

Activated carbon can be used to remove many differ-ent pesticides, pharmaceuticals, and estrogenic com-pounds (Robeck et al., 1965; Steiner and Singley, 1979;Miltner et al, 1989; Pirbazari et al, 1992; Sacher et al.,2000; West, 2000). The performance of activated carbondepends on the properties of the activated carbon sorbent(surface area, pore size distribution, surface charge, oxy-gen content) and on the properties of the solute (shape,size, charge, and hydrophobicity). Hydrophobic interac-tions are the dominant mechanism of removal for mostorganic compounds in activated carbon adsorption sys-tems. However, ion exchange interactions can result inremoval of polar solutes (Youssef et al., 1982; Mat-sumura et al., 1985; Crittenden et al., 1999). As a resultof the hydrophobic interactions, activated carbon effi-ciently removes most nonpolar organic compounds (i.e.,those compounds with log KOW . 2). The ability of ac-tivated carbon to remove more polar compounds will de-



pend upon the strength of the polar interactions, whichare difficult to predict a priori.

NOM in water competes for adsorption sites and de-creases the activated carbon capacity for micropollutants(Newcombe et al., 1997a, 1997b; Pendleton et al., 1997;Newcombe, 1999; Wu and Pendleton, 2001). As a result,treatability studies conducted in distilled water should beinterpreted with caution because they do not considercompetition from NOM. For example, addition of 10 to20 mg/L of powdered activated carbon (PAC) to distilledand river water spiked with seven antibiotics achievedbetween 50% and greater than 99% removal. However,when the same experiments were repeated in river watercontaining 10.7 mg/L of NOM, removal decreased by 10to 20% (Adams et al., 2002).

Although results of full-scale treatment studies withEDCs and PPCPs are not available, considerable infor-mation is available on the ability of activated carbon toremove compounds that cause taste and odor problems.PAC is usually added in presedimentation or contactbasins prior to coagulation, sedimentation, and filtration.Usually, 1 to 3 h of contact is provided for the PAC, af-ter which the PAC settles out in the sedimentation tank,and is then disposed with other WTP sludges. For ex-ample, PAC added at dosages ranging from 5 to 50 mg/Lremoved greater than 90% of methylisoborneol (LogKOW 5 3.1) in raw water (Gillogly et al., 1998; Zhangand Emary, 1999; Bruce et al., 2002). Under the condi-tions encountered in drinking water treatment plants, re-moval of micropollutants by PAC tends to be indepen-dent of initial contaminant concentrations (Knappe et al.,1998; Leung and Segar, 1999). In a laboratory, 20 mg/Lof PAC with 1-h contact time reduced the concentrationof the moderately hydrophobic pesticide lindane (LogKOW 5 3.72) from 10 to 0.1 mg/L (Kouras et al., 1998).Laboratory studies conducted with EDCs suggest thatPAC will be effective at removing between approxi-mately 60 and 80% of nonylphenol and NPEs (Carlile,et al., 1996; Fielding et al., 1998). Unpublished PAC ex-periments in distilled water and/or surface water con-ducted by the authors suggest similar removals of E2,EE2, triclosan, dilantin, bisphenol A, and octylphenol.

Granular activated carbon (GAC) systems operate asadsorptive packed beds or filters. While PAC is added,contacted with water (,4 h), and removed (settling/fil-tering), GAC systems operate with stationary phase(GAC packed beds), continuous water flow, and contacttimes of less than 30 min. GAC systems remain in oper-ation for months to years, and thus achieve pseudo-equi-librium with micropollutants in the influent water (Crit-tenden et al., 1991). The influent micropollutant andDOC concentrations, contact time with the GAC, and

type of GAC impact the adsorption and breakthrough ofthe micropollutant across GAC systems. For example, theadsorption characteristics of three types of activated car-bon for radiolabeled 17b-estradiol were studied to deter-mine the time necessary to reach equilibrium between thesolid and the liquid phase (Fuerhacker et al., 2001). 17b-Estradiol was quickly adsorbed and conditions close toequilibrium were reached after 50–180 min. The equi-librium concentrations were calculated to be at 49–81%of the initial concentrations between 1 and 100 ng/L, with0.51 ng/L remaining for a 1 ng/L initial concentration andbetween 5.9 and 14.6 ng/L for a 100 ng/L initial con-centration. Wang and Lee (1997) determined GAC (coalbase) design criteria to remove various micropollutantsincluding organophosphorus, volatile organic chemicals,and phenol found in drinking water sources having dif-ferent background DOC concentrations. It was found thatsorption capacities of the micropollutants were reducedwith increasing background DOC concentration due pre-sumably to DOC pore blockage of the GAC and reduc-tion of the surface area available for the adsorption ofmicropollutants. The results obtained from activated car-bon filtration for the removal of nonylphenol showed thatcontact times of 4 days and 24 h and activated carbondosages of 0.1 and 1 g/L could be obtained withnonylphenol total contaminant loadings up to 0.01 mg/gon saturation of the GAC (Tanghe and Verstraete, 2001).The influence of temperature (4 or 28°C) on nonylphe-nol adsorption on GAC was negligible and the sorptioncapacity of dissolved humic acids for nonylphenol wasconsiderable at a nonylphenol concentration of 10 mg/L.

The use of GAC leads to very high removal of mi-cropollutants during the first weeks/months, but overtime, more strongly adsorbable constituents can displacepreviously adsorbed compounds. The relative strength ofadsorption is often characterized by Freundlich isothermparameters (e.g., sorption capacity, K ). Freundlich “K”values are available for numerous compounds regulatedunder the Safe Drinking Water Act (Speth and Adams,1993; Speth and Miltner, 1998), and correlated with phys-ical properties of organic compounds (Crittenden et al.,1999). Freundlich constants can be used to predict PACand GAC performance (Knappe et al., 1998; Najm et al.,1991a, 1991b). Typically, when K values were greaterthan 200 the activated carbon process is considered tech-nically and economically feasible. Although K values arenot currently available for emerging EDCs and PPCPs,quantitative structure–activity relationship models couldbe used to estimate the values. Generally, compoundswith log KOW values . 2 have Freundlich K values .

200 and, therefore, many EDCs and PPCPs (Table 1) maybe amenable to removal by PAC or GAC.

458 SNYDER ET AL.

Oxidation processes

In water treatment systems, several of the commonlyused disinfectants also can result in transformation ofEDCs and PPCPs. In drinking water treatment systems,chlorine, chlorine dioxide, and ozone are frequently usedwhile disinfection of wastewater effluent usually is lim-ited to chlorine. Oxidation of EDCs and PPCPs is selec-tive for certain chemical structures and functional groups.Among the three oxidants, ozone tends to be the most re-active; however, all three oxidants are strong elec-trophiles that exhibit similar trends of reactivity with or-ganic compounds. As a result, certain generalizations canbe observed in the reactivity of the oxidants (Hoigne andBader, 1983a, 1983b, 1994; Hoigne et al., 1985; Larsonand Weber, 1994; Tratnyek and Hoigne, 1994; Gallardand von Gunten, 2002): (1) dissociated acidic compoundsare more reactive than protonated forms (i.e., reactivityincreases with pH), but nondissociated bases are more re-active when not protonated; (2) general order of reactiv-ity from highest to lowest for aromatic or aliphatic com-pounds: thiols . amines . hydroxyl . carboxyl; and (3)aromatic compounds are more reactive than aliphaticcompounds.

Chlorination. In the United States, free chlorine (i.e.,HOCl and OCl2) is commonly used for disinfection andoxidation of reduced inorganic species such as Fe(II),Mn(II), and S(-II). In the presence of ammonia, free chlo-rine will produce chloramines, which tend to be less re-active than free chlorine (Weil and Morris, 1974). Forexample, the application of free chlorine resulted in theremoval of .95% of the herbicide glyphosate in OhioRiver water after 15 min while monochloramine did notremove any of the herbicide under similar conditions(Speth, 1993). The reduced reactivity of chloramine isparticularly important in wastewater treatment plants be-cause enough ammonia will be present to convert freechlorine into monochloramine in wastewater effluent thathas not undergone nitrification.

Free chlorine reacts rapidly with phenolic compounds,mainly through the reaction between hypochlorous acidand the deprotonated phenolate anion (Faust and Hunter,1967). The reaction results in sequential chlorine addi-tion to the aromatic ring followed by ring cleavage. Thereactivity of the phenolic functional group likely explainsthe transformation of estrogenic hormones and nonylphe-nol by chlorine observed in laboratory studies (West,2000). However, these results conflict with findings inJapan that report chlorine as ineffective for the degrada-tion of estrogens (Matsui and Takigami, 2000).

The transformation of several amine-containing anti-biotics (Adams and Kuzhikannil, 2000; Huang et al.,

2001), diclofenac (Sedlak and Pinkston, 2001), and caf-feine (Gould and Richards, 1984), was observed in lab-oratory experiments with chlorine. Additional studies areneeded to assess the reactivity of compounds with otherpotentially reactive functional groups and the relative im-portance of these reactions under conditions used for dis-infection of water and wastewater. In addition, the reac-tions between chlorine and EDCs and PPCPs necessitatequenching of chlorine in samples that may have a chlo-rine residual, especially when 24-h composite samplesare collected. For example, chlorination (1 mg/L; pH 5

7.5) removed .90% of seven amine-substituted antibi-otics from river water within 40 min (Adams andKuzhikannil, 2000). Therefore, careful attention must bedirected at discerning removal by chlorine vs. removalfrom a quenching agent.

Chlorine dioxide. Chlorine dioxide (ClO2) can oxidizeherbicides, pesticides, and PAHs, and is generally astronger and faster oxidant than free chlorine (Ravachaand Blits, 1985). Chlorine dioxide application (1–1.5mg/L for 10 min) at two full-scale WTPs removed lessthan 30% of the total pesticide concentration (0.05 to 3.5mg/L) in the raw water (Griffini et al., 1999). ClO2 wasmore effective than free chlorine at oxidizing amine-con-taining pesticides (phenylamide) and herbicides (ame-tryn) than for other herbicides (isoproturon, glyphosate)(Eldib and Aly, 1977; Speth, 1993; Lopez et al., 1997).Chlorine dioxide (2 mg/L at pH 7) removed 50% of sev-eral PAHs, but contact times required for this level of ox-idation ranged between 6 s (benzo[a]pyrene) to 17 h(benzo[b]fluranthene) for different PAHs (Ravacha andBlits, 1985). No additional information is available onoxidation of EDCs or PPCPs by ClO2. However, we an-ticipate that ClO2 will react with compounds containingphenolic amino and thiol functional groups (Hoigne andBader, 1994).

Ozonation. Ozone (O3) is used in water treatment asboth a disinfectant and an oxidant. During ozonation, twostrong oxidants can lead to transformation of EDCs andPPCPs and other organic compounds: molecular O3 andhydroxyl radicals (HO�) (Hoigne and Bader, 1983a,1983b). O3 is a selective electrophile that reacts withamines, phenols, and double bonds in aliphatic com-pounds, while HO� reacts less selectively with second-order rate constants (kHO) on the order of 108 to 1010

M21s21 (Buxton et al., 1988; Haag and Yao, 1992). Asa result of the selective nature of ozone, transformationof micropollutants may require the use of advanced oxi-dation processes (AOPs), such as UV/hydrogen peroxide(H2O2), O3/H2O2, or UV/O3 (Hoigne, 1998; Acero et al.,



2000; Acero and Von Gunten, 2001). Given informationon water quality (e.g., alkalinity, pH, and NOM) it maybe possible to estimate the concentrations of O3 and OHpresent during ozone treatment. Such data can be usedwith second-order rate constants to predict the rate oftransformation of EDCs and PPCPs during ozone treat-ment.

Preliminary data for a set of EDC/PPCP compounds in-dicate a wide range of reactivity for reactions with ozone;roxithromycin (4.5 3 106 M21s21), diclofenac (.105

M21s21), ethynylestradiol .105 M21s21), carba-mazepene (0.78–3 3 105 M21s21), sulfamethoxazole(105 M21s21), bezafibrate (590 M21s21), ibuprofen (6M21s21), diazepam (0.95 M21s21), and iopramide (,0.8M21s21) (Huber et al., 2002). Most kHO second-order rateconstant values were on the order of 5 3 109 M21s21.Under conditions encountered in water treatment systems([HO�]/[O3] < 1028), only those compounds with ozonerate constants (kO3) greater than ,50 M21s21 will betransformed to an appreciable degree (.50%) through di-rect reactions with ozone. Therefore, during ozonation themechanism of oxidation will be O3 or a combination ofO3 and HO�. Some reports have documented the removalof EDCs and PPCPs in bench-scale treatment systems. Forexample, ozonation removed diclofenac, carbamazepine,and bezafibrate, but not clofibric acid (Sacher et al., 2000).Improved removal of clofibric acid, ibuprofen, and di-clofenac occurred when ozonation was conducted in thepresence of hydrogen peroxide (0.4 to 0.7 mg H2O2/mgozone dosed) (Carlson et al., 2000). Estrogen steroids andnonylphenols also reacted with ozone under conditionscomparable to those encountered in water treatment sys-tems (West, 2000). Ozonation of estrogenic chemicals inspiked groundwater indicated some removal of mestranol,estradiol, ethynylestradiol, norethistrone, ethistrone, andestriol by aeration alone, and higher removals with ap-plied ozone (Carlile et al., 1996). However, most EDCsand PPCPs have low Henry’s Law constants, and aera-tion would not appear to be an effective means for re-moval. In separate work, a linear relationship betweenozone dose and fraction of E2 remaining in model waterswas apparently observed during O3 and O3/H2O2 pro-cesses (Kosaka et al., 2000).

Ultraviolet (UV) irradiation. UV lamps are usedwidely for microbial disinfection of water and waste-water. In several cases, they also have been used for treat-ment of micropollutants (Mofidi et al., 2000). Becauseseveral EDCs and PPCPs have chromophores that leadto adsorption of light at UV wavelengths, many may beamenable to transformation during UV treatment. How-ever, typical UV doses required for disinfection (i.e., ,5to 30 mJ/cm2) are several orders of magnitude lower than

those used for treatment of micropollutants. Therefore,UV treatment of EDCs and PPCPs probably will not beeconomically competitive with other advanced treatmentmethods (e.g., reverse osmosis). As discussed previously,UV treatment in combination with ozone or hydrogenperoxide may be practical in some situations.

UV alone or as part of AOP systems has been used tooxidize pesticides and other micropollutants, primarily ingroundwater (Beltran et al., 1992, 1993, 1996, 2000; Lar-son and Weber, 1994; Chiron et al., 2000; De Laat et al.,1999). Few reports of EDC or PPCP removal during UVtreatment exist. For example, oxidation of fragrances (ni-tromusks) is achieved using UV light produced by low-pressure mercury lamps with faster transformation rateswith the addition of hydrogen peroxide (Neamtu et al.,2000). At UV doses of 3000 mJ/cm2, the removal of dif-ferent antibiotics from distilled and natural waters rangedbetween 50 to 80% for the seven antibiotics, which wereall selected because the compounds were detectable byHPLC.

Biotransformation

Biotransformation may provide a basis for the cost-ef-fective removal of EDCs and PPCPs from water. Avail-able data from municipal wastewater treatment systemssuggest that many of the compounds are transformed bybacteria (Johnson et al., 2000; Johnson and Sumpter,2001; Ternes et al., 1999a, 1999b). However, little isknown about the importance of biotransformation indrinking water treatment plants or in soil. Data from siteswhere wastewater effluent is used to recharge aquifers(i.e., soil aquifer treatment systems) indicate that manyof the PPCPs are removed during the first few weeks ofpassage through the aquifer (Drewes et al., 2001b). How-ever, certain recalcitrant compounds including theantiepileptics, carbamazepine and primidone, and iodi-nated X-ray contrast media persist during infiltrationthrough vadose and saturated soil zones (Drewes et al.,2001a, 2001b). Similar results were observed in bank fil-tration (Heberer, 2002) and during slow sand filtration indrinking water treatment plants (Sacher et al., 2000).These studies confirm that some PPCPs are biodegrad-able, while others are not or have not yet been investi-gated.

Membrane separation

Most organic EDC/PPCP compounds range from 150to 500 Daltons in molecular size. As a result, only thosecompounds associated with particles or colloidal organicmatter will be removed during microfiltration and ultra-filtration (UF). For example, microfiltration (0.1 to 0.4mm) did not remove two steroid hormones and six dif-

460 SNYDER ET AL.

ferent acidic drugs and beta blockers at two different full-scale water recycling systems treating tertiary municipalwastewater effluent (Huang and Sedlak, 2001). MostEDCs and PPCPs will be removed by reverse osmosis(RO) and tight nanofiltration (NF) systems (i.e., thosewith a low molecular weight cutoff). For example, ROcan achieve greater than 90% removal of steroid hor-mones (Huang and Sedlak, 2001).

Experience with other organic/inorganic compoundssuggests that polar compounds and charged compoundsthat interact with membrane surfaces will be better re-moved than less polar or neutral compounds (Huxstepand Sorg, 1988; Duranceau et al., 1992; Berg et al., 1997;Carlson, 2000; Yoon et al., 2001, 2002a, 2002b). For ex-ample, the removal efficiency of several low molecularweight compounds (,150 Daltons) increased at higherpH due to electrostatic repulsion between RO membraneand dissociated organic compounds (Ozaki and Li, 2002).In the same work, removal of neutral organic compoundsincreased linearly with molecular weight and molecularwidth. Hydrophobicity of the micropollutant and mem-brane also affect removal. For example, pesticide ad-sorption on, and removal by, RO and NF membranes wascorrelated with Log KOW (Kiso et al., 2001a, 2001b).

Several recent laboratory studies also have demon-strated the removal of PPCPs from water. Compoundsremoved by RO included several different antibiotics(Adams et al., 2002). Results from laboratory studies con-ducted by adding compounds to deionized water shouldbe interpreted with caution because the presence ofcations and NOM can alter removal efficiency. RO, NF,and charged UF membranes can remove inorganic EDCs(e.g., arsenic, perchlorate) (Uludag et al., 1997; Urase etal., 1997; Chianese et al., 1999; Vrijenhoek and Waypa2000; Van der Bruggen et al., 2001; Yoon et al., 2002a).Overall, membrane separation provides an excellent bar-rier for most EDCs and PPCPs, except the lower molec-ular weight uncharged compounds.

SUMMARY OF WATER TREATMENTPROCESSES FOR EDC AND PPCP

CONTROL

Table 2 provides a generalized summary of the poten-tial for removal of various classes of EDCs and PPCPs.Estimates of removal are based upon chemical structure(size, hydrophobicity, functional group composition).

1. Coagulation would only be expected to remove hy-drophobic compounds associated with particular orcolloidal material with high organic carbon content.

2. Activated carbon adsorption removes hydrophobic

compounds very well, but competition between morepolar or larger molecular weight compounds and othermatrix organics has not been well documented.

3. Oxidation will preferentially attack compounds withelectron-activating functional groups (thiols, amines,hydroxyl) located near C C bonds (benzene rings).Organic compounds will react more rapidly withozone than chlorine dioxide or chlorine.

4. Membranes provide a physical barrier capable of highremovals, but removals are dependent upon compoundstructure (size, polarity) and membrane properties.

Given experience with removing pesticides and othertrace micropollutants, water treatment technologies canbe integrated into existing WTPs or designed into newfacilities to assure high levels of EDC and PPCP com-pound removal. For example, simple process modifica-tions such as addition of powder-activated carbon to aconvention WTP during periods of high-risk of EDC orPPCP in the raw water (e.g., low streamflow during thesummer comprised primarily of upstream wastewater dis-charge) could provide high levels of removal. Incorpo-ration of more advanced technologies could require ex-tensive modifications to WTP operations. For example,inclusion of O3 or AOP systems should be followed bybiofiltration to remove polar oxidation byproducts. Com-binations of treatments may be required to remove mix-tures of compounds that may occur in a raw drinking wa-ter (e.g., membranes, oxidation).

Future research needs

A great deal of additional data are needed to under-stand the relevance of trace EDCs and PPCPs in waterand how these compounds may be removed by watertreatment. It is important to gain more information as tothe toxicological impacts of trace levels of EDCs andPPCPs in water. Once these impacts are quantified, safeexposure limits can be established that will allow the wa-ter industry to determine “good removal” rates. Stan-dardized analytical methods for detection of commonlyoccurring EDCs and PPCPs are critical. These methodsshould be based on equipment that most laboratoriescould afford and have expertise to operate. Once analyt-ical methodologies are available, environmental screen-ing should include testing for bioaccumulation of EDCsin wildlife and humans. Future studies should seek toidentify population-level impacts of EDCs and PPCPs onwildlife and relate these effects to biomarkers.

Although reports of EDC and PPCP removal by wa-ter treatment are beginning to become available, much isstill unknown as to the fate of these emerging contami-nants in WTPs. It is critical to understand the size dis-



tribution of EDCs and PPCPs and their association withparticulate and colloidal materials in raw drinking watersupplies to assess removal across conventional WTPs de-signed for particular and colloidal removal. The oxida-tion capability of chloramines on EDC/PPCP compoundsreactive with chlorine should be quantified. Oxidationbyproducts should be identified and tested for both en-docrine and carcinogenic health risks. Effects of nonionicpolymers on improving EDC/PPCP removal during coagulation have not been reported. Aerobic biofilmsshould be compared to nonbiological sand filters to quan-tify the potential benefits of changing the point of chlo-rination in a conventional WTP to optimize EDC/PPCPremoval. Breakthrough curves should be generated forGAC packed columns with a variety of EDC and PPCPcompounds in water containing NOM to evaluate poten-tial displacement/desorption of organics over time. Par-tition and permeation coefficients across membranes forEDCs and PPCPs should be determined. Additionally,membranes should be evaluated for leaching of plasti-cizers that may act as EDCs. Integrated water manage-ment options, which include source water protection,should complement advanced drinking water treatment.

ACKNOWLEDGMENTS

Funding for this review was provided in part by theAmerican Water Works Association Research Founda-tion (AWWARF) project 2758 entitled “Evaluation ofConventional and Advanced Treatment Processes to Re-move Endocrine Disruptors and Pharmaceutically ActiveCompounds.”

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